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Seasonal distribution, gas–particle partitioning and inhalation exposure of brominated flame retardants (BFRs) in gas and particle phases

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Abstract

Background

Atmosphere is one of the sources and sinks to gas- and particle-bound brominated flame retardants (BFRs). Therefore, BFRs can enter human body via inhalation. In the present study, 79 of gas- and particle-phase samples (TSP, PM10 and PM2.5) were collected during 2015–2016 in urban area of Shanghai, China to investigate the occurrence of 25 polybrominated diphenyl ether (PBDEs) congeners, hexabromobenzene (HBB) and 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE).

Results

The total concentrations of BFRs ranged from 0.66 to 13.7 pg/m3, 25.82 to 376.27 pg/m3, 14.58 to 365.49 pg/m3 and 15.17 to 304.89 pg/m3 in gas, TSP, PM10 and PM2.5, respectively. HBB was the main compound in gas phase; while BDE-209 was the dominant congener in particle phase. Atmospheric BFRs in winter was much lower than in summer, while particle phase showed opposite seasonal tendency.

Conclusions

Clausius–Clapeyron equation suggested that BTBPE, BDE-28, -66, -99, -100, and -154 were more driven by temperature and evaporated from local contamination. Higher-brominated congeners tended to be absorbed in particle phase, while low-brominated compounds partitioned in both gas and particle phases. Gas/particle partitioning results showed that absorption into the aerosol organic matter was the dominant process for BFRs. The inhalation health risk assessment demonstrated that the hazard quotient (HQs) for most PBDEs in winter was higher than in summer.

Background

Brominated flame retardants (BFRs) are substances added to consumer products to prevent from firing and are widely used in textiles, furniture, electronics, vehicles and other materials [1]. It has been proved that BFRs are ubiquitous in all kinds of environment matrices [1, 2] and can be detected in biology and human [3, 4]. Studies suggested that BFRs could interfere with the reproductive and endocrine system, and cause thyroid disorders, neurobehavioral disorders and other health problems [5, 6].

Polybrominated diphenyl ethers (PBDEs) are the most commonly used compounds among BFRs for decades. Penta-BDE, Octa-BDE, and Deca-BDE are three major technical PBDE products used in the market. Due to the high environmental persistence, bioaccumulation tendency and toxicity [1], they were added to the list of persistent organic pollutants (POPs) and have been phased out gradually [7, 8]. With the extensive restrictions on the use of PBDEs, novel brominated flame retardants (NBFRs) such as hexabromobenzene (HBB), 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE) have been used to meet flammability standards. HBB is used in woods, textiles, electronics and plastics with production volume of 600 t in Qingdao, China annually [9]. BTBPE, currently, is a replacement for the Octa-BDE mixtures and used in polymeric materials, cotton and polyester [9, 10].

As typical semi-volatile organic compounds (SVOCs), atmosphere plays an important role in transportation and transformation of BFRs. The detection of high levels of these compounds in Antarctica and Arctic sites [11, 12] suggests that long-range transport process occurs in atmosphere. It can be accepted that temperature and particle size are important factors influencing the environmental fate of these compounds [13,14,15]. For example, the partial pressure in gas phase and gas–particle partitioning of most low-molecular-weight PBDEs (except for BDE-85), HBB and BTBPE were significantly correlated with the ambient temperature [16]. Another research found that these compounds tended to be enriched in particles < 50 µm for road dust and hardly detected in particles > 50 µm [13].

Although numerous research have been conducted BFRs in atmosphere worldwide [17,18,19], only few studies investigated the mechanisms that influence their transport and gas–particle partitioning [20, 21]. Furthermore, there is limited information regarding the seasonal distribution of BFRs and the correlations among BFR levels, particle size and temperature change. In the present study, we examined a total of 24 PBDEs and 2 NBFRs in gas and particle samples from July 2015 to January 2016 in Shanghai, China. The objectives of this project were: (a) to investigate the concentration and composition of BFRs; (b) to assess the seasonal variation and temperature effect of BFRs; (c) to demonstrate the distribution mechanism of gas–particle partitioning for BFRs, and (d) to estimate the human inhalation exposure of BFRs-contaminated air.

Materials and methods

Chemicals and materials

Mixed standards for 9 PBDE congeners (including BDE-28, -47, -66, -85, -99, -100, -153, -154, -183) and for other 15 PBDE congeners (including BDE-194, -195, -196, -197, -198, -199, -200, -201, -202, -203, -204, -205, -206, -207, -209) were both obtained from Wellington Laboratories. These 24 PBDE congeners can be divided into 8 homologues (Additional file 1: Table S1): Tri-BDE, Tetra-BDE, Penta-BDE, Hexa-BDE, Hepta-BDE, Octa-BDE, Nona-BDE and Deca-BDE. Two NBFRs including HBB and BTBPE, the surrogate standards (BDE-77, BDE-138) and internal standard (BDE-118) were purchased from Accu Standard (New Haven, CT, USA). All of the organic solvents were in pesticide residue grade and obtained from CNW Technologies GmbH. Sulfuric acid was a guaranteed reagent (Sinopharm Chemical Reagent Corporation, Beijing, China). Silica gel (0.063–0.200 mm) was of chromatographic grade (Merck, Darmstadt, Germany) for chromatography.

Sample collection

A total of 79 samples [14 polyurethane foam (PUF), 26 total suspended particulate (TSP), 20 particulate matter with particle size below 10 µm (PM10), 19 particulate matter with particle size below 2.5 µm (PM2.5)] were collected from July 2015 to January 2016 in a typical urban region of Shanghai, China. The sampling site is located in the northwest of Shanghai (N31°14′47.35″, E121°21′50.80″), mounted on roof approximately 15 m above the ground. A highway is located in the southeast of sampling site at around 150 m. 1 km to the north was Beijing–Shanghai high-speed railway. Others were residential area around.

Samples were carried out using two high-volume air samplers (ECHO HiVol, Italy) at 24 h per sample. Gas and particle phases were isolated by drawing air through PUF (65 mm × 75 mm, TISCH, USA) and quartz fiber filter (QFF, 102 mm in diameter, Munktell, Sweden) at a flow rate of about 225 L/min. The sampled air volume was approximately 300 m3. PM2.5 and PM10 were collected using PM2.5 and PM10 cyclones, respectively.

Prior to sampling, PUF plugs were Soxhlet extracted for 48 h with acetone/ cyclohexane (1:1, v:v), while QFFs were baked at 400 °C for 5 h and wrapped with aluminum foils to remove any residue organic pollutants. After sampling, the PUFs and QFFs were wrapped in aluminum foils, sealed in polyethylene zip bags separately, and then transported to laboratory and stored at − 20 °C until extraction. Unloaded and loaded QFFs were weighed after 24-h conditioning at constant temperature and humidity chamber to determine the concentrations of TSP, PM10 and PM2.5.

Sample pretreatment

Pretreatment procedures followed the previously published method [4] with minor modifications. Briefly, each PUF and QFF sample was spiked with BDE-77 and BDE-138 (2 ng) and then Soxhlet extracted for 48 h with 150 mL acetone/cyclohexane (1:1, v:v) at 65 °C. After evaporated to about 3 mL by a rotary vacuum evaporator (R-210/215, BUCHI Labortechnik), 3 mL concentrated sulfuric acid was added to remove lipids and organic matters. A modified silica column (0.1 g silica + 0.9 g acidic silica + glass fibers at bottom) was used for further clean up. The extract volume was evaporated to 200 μL under a gentle stream of nitrogen and then BDE-118 (2 ng) was added before instrumental analysis.

Instrumental analysis

The analysis of BFRs was also carried out based on our previous method [4]. PBDEs and NBFRs were quantified with Agilent 6890 gas chromatograph/5975 mass spectrometer connected with a DB-5MS column (15 m × 0.25 mm × 0.10 µm, J&W Scientific). The negative chemical ionization (NCI) source in selected ion monitoring (SIM) mode was employed. All the target compounds were quantified by scanning bromide ion fragments (m/z 79 and 81). Helium was used as carrier gas to carry the extracts at a set constant flow of 1.4 mL/min, and methane was used as the chemical ionization moderating gas. The ion source, injector, and transfer line temperature was set at 200 °C, 280 °C and 290 °C, respectively. The GC oven temperature program was given as followed: 80 °C for 1 min, 15 °C/min to 300 °C, 2 °C/min to 310 °C and held for 5 min. Each sample was injected in the splitless mode with the injector volume at 1 μL.

Quality assurance/quality control (QA/QC)

Before sample collection, gas-phase breakthrough test was conducted using a second PUF plug in series with the first one. The concentrations of BFRs in the second plug were less than 10% compared to the first plug which showed that PUF breakthrough could be ignored. Procedural blanks were processed for every season while laboratory blanks were taken for every month during the pretreatment procedure to check background contamination. Drift blanks were run after every 10 samples, and the results were acceptable when the relative standard deviation for individual PBDE congener in standards was less than 10%. The limit of detection (LOD) and the limit of quantification (LOQ) were defined as a signal of three times and ten times the noise level, respectively. More details about LOD and LOQ are provided in Additional file 1: Table S1. BFRs concentrations in all samples were blank corrected. Recoveries of BDE-77 and BDE-138 ranged from 60 to 110% and 78 to 115%, respectively. The reported concentrations in this study were not corrected with surrogate recovery.

Statistical analysis

SPSS Version 20 was used for Spearman correlation analysis to investigate correlations between data and other statistical analysis. The normality test suggested non-normal distribution of data, and therefore statistics were expressed as geometric mean (GM), geometric standard deviation (GSD) and median value. Mann–Whitney Test was performed to assess the significance of differences. The level of significance was assigned at 0.05.

Results and discussion

Concentrations and distributions of TSP, PM10 and PM2.5

Particle concentrations of TSP, PM10 and PM2.5 were ranged from 57.5 to 274.4 μg/m3, 36.7 to 199 μg/m3 and 30.3 to 183.2 μg/m3 (Additional file 1: Fig. S1a), which were comparable to those reported in China and other Asia countries [22,23,24,25,26,27,28], but much higher than developed countries in the UK, Spain and Greece [29] (Additional file 1: Table S2). The median contribution of particles between 0–2.5 µm, 2.5–10 µm and 10–100 µm was 59.7%, 13.4% and 26.9%, respectively (Additional file 1: Fig. S1b), suggesting that finer particles are the dominant fraction in TSP.

Strongly negative correlations were found between temperature (except continued high temperature weather in Jul. and Aug.) and TSP (P < 0.05), PM10 (P < 0.01), PM2.5 (P < 0.01) (Additional file 1: Fig. S2). Previous studies [30, 31] reported similar temperature pattern in PM2.5. This is probably because higher temperature can accelerate air convection, which brings about the dilution and diffusion of the particles [31, 32]. Besides, increased emission by heating and power production in winter can cause enhanced particle concentrations in China.

Concentration of BFRs

The range and geomean concentrations of gas and particle phases are given in Table 1. Low-brominated compounds (Tri-, Tetra-, Penta-, Hexa-, Hepta- BDE and HBB) and BTBPE had higher detected frequency as 85–100% in gas phase. All the BFR compounds were detected in all particle phase (100%) except BDE-28, -47, -66, -99, -100, -183 and -202 with the detection frequency ranged from 25 to 94%. The total concentrations of BFRs (Σ25BFRs) ranged from 0.66 to 13.7 pg/m3, 25.82 to 376.27 pg/m3, 14.58 to 365.49 pg/m3 and 15.17 to 304.89 pg/m3 in gas, TSP, PM10 and PM2.5 and the GM concentrations were 5.96, 101.70, 49.50 and 44.21 pg/m3, respectively.

Table 1 Concentrations of BFRs in gas and particle phases (pg/m3)

HBB and BDE-47 were the most abundant congeners in atmosphere with the GM concentration 1.26 and 1.03 pg/m3, followed by BDE-28 and -66. In particle phase, BDE-209 had the highest GM concentration as 62.56, 29.01 and 21.53 pg/m3 in TSP, PM10 and PM2.5, respectively. BDE-207 was secondly abundant in particle phase, followed by BDE-206 and -208.

In gas phase, the GM level of Σ7PBDEs (-28, -47, -99, -100, 153, -154, -183) from this study was 5.8 and 0.9 pg/m3 in summer and winter (Additional file 1: Table S3), which was comparable with those in Yangtze River Delta, China [33], and higher than some European countries like Norway, France [34], but much lower than some other Asian countries like Kuwait [35] and Philippines [34]. Compared with America, the concentration in present study was slightly lower than in St. Lawrence Island and Point Reyes [34]. Besides, it was significantly lower than the research from Guiyu which is a typical e-waste recycling site in China [14] with the Σ7PBDEs 4073.89 and 1916.5 pg/m3 in summer and winter, respectively. The atmospheric GM level of HBB in this study (2.5 pg/m3 in summer) was in accordance with those in Beijing, Xiamen, Tianjin and slightly lower than in Guangzhou [36] in China.

Compared with the PBDEs in TSP from other studies, the GM concentration of Σ7PBDEs in this study (2.20 and 14.37 pg/m3 in summer and winter) was similar to those in Kuwait [35], Italy [37], and significantly lower than those in Guiyu with the average Σ7PBDEs concentration as 2068 and 15,229 pg/m3 in summer and winter [14]. The level of BDE-209 in TSP was 47.79 and 59.36 pg/m3 in summer and winter, which was higher than in Africa [38], France [38], Italy [37], but much lower than in Shanghai [39], Harbin [40] and Guiyu [14] in China. Additionally, the levels of Σ7PBDEs in PM2.5 in this study were slightly higher than in East China Sea [41] and Spain [42]. Overall, the GM levels of gaseous PBDEs in present study were higher than in European countries and comparable to other cities in China, but lower than in America and other Asian countries. In particle phase, the concentration of PBDEs in this study was moderate compared with other cities and countries. However, all the results in this study were significantly lower than in the e-waste city Guiyu in China.

Congener profiles of BFRs

Compositions of BFRs in both gas and particle phases are shown in Additional file 1: Fig. S3. The predominant BFR congeners in gas phase were HBB, Tetra-BDE (BDE-47, BDE-66) and Tri-BDE (BDE-28), accounting for 39%, 25% and 14% among the total BFRs. The relatively high abundance of these lighter BFRs was mainly due to their higher vapor pressure [43]. Similarly, BDE-28, -47 were found to be the major atmospheric congeners in an earlier research [44] in Shanghai rural regions, which accounted for 13.1% and 11.7% in average. Deca-BDE (BDE-209) was the dominant congener in particle phase contributing to more than half of the Σ25BFRs. The excessive of BDE-209 indicated that commercial Deca-BDE may still be in use in the studied area. This result was consistent with another research which reported that BDE-209 contributed 65% and 59% of gaseous plus particle-bound phase at urban and rural sites, respectively, in Lake Chaohu, China [21]. Nona-BDE (BDE-206,-207,-208) was secondly abundant with 17%, 21% and 23% in TSP, PM10 and PM2.5.

The ratios of BDE-47/99, BDE-99/100 are used to illustrate the source and transformation of PBDEs. In this study, the average ratio of BDE-47/99 in total air concentrations (gas plus particle phase) was 1.04, which was in accordance with other researches in the Arctic (1.1) [45], and slightly lower than those in Lake Chaohu, China (2.6) [21], the USA (1.4–2.4) [46]. This ratio was similar to those in commercial Penta-BDE mixture DE-71 (1.27) and Bromkal 70-5DE (1.05) [47], suggesting the very likely use of these technical products in the sampling site. The ratio of BDE-99/100 in this study (1.92) was comparable to the ratios in Lake Chaohu, China (1.4) [21], northern China (2.4) [48], and lower than other researches reported in the Arctic (5.4) [45], the USA (4.1–5.1) [46], the commercial DE-71 (3.71) and Bromkal 70-5DE (5.73) [49]. For outdoor environment, photolysis under sunlight and photooxidation by ·OH radicals occurs in PBDEs with different degradation rates for each compounds [15, 50]. BDE-47 and -100 show higher degradation rates than BDE-99; thus, ratios of BDE-47/99 and BDE-99/100 are estimated to be lower and higher, respectively [21]. However, higher-brominated PBDE congeners (such as BDE-209) could also degrade into BDE-47, -99 and -100, resulting in the variation of these ratios [38, 44, 51]. Therefore, low ratio of BDE-99/100 in this study was more likely affected by the degradation of high-brominated compounds instead of the different degradation rates.

BTBPE showed significant correlations with most of the BFR homologues and the highest coefficient were found in Hexa-, Hepta- and Octa-BDE (Additional file 1: Tables S4–S7). This was consistent with prior observations as BTBPE is a replacement for commercial Octa-BDE mixtures [52, 53]. Higher correlations were observed among homologues with similar number of bromine, such as Tetra-, Penta- and Hexa-BDE (r = 0.675–0.955, P < 0.01), or Octa-, Nona-, and Deca- BDE (r = 0.562–0.873, P < 0.01). Weaker or no significant correlations were found between low-brominated and high-brominated congeners (Octa-, Nona-, Deca-BDE). It was possibly due to complicated environmental fates (e.g., orders magnitude differences of octanol–air partition coefficients) and different usage of commercial BDE mixtures [52].

Seasonal variations and temperature effect of BFRs

The levels of NBFRs and low-brominated congeners in gas phase were ranked as summer > autumn > winter (Additional file 1: Fig. S4). Different tendency was observed in particle phase: the concentration of NBFRs, lower-brominated congeners and Octa-BDE were in the order as winter > autumn > summer; whereas, higher-brominated congeners (Nona- and Deca-BDE) were as autumn > winter > summer.

Low-brominated congeners in gas phase during summer were much higher than in winter with GM concentrations as 8.41 and 1.98 pg/m3, respectively (Fig. 1). Other researchers also observed elevated atmospheric concentration of BFRs in summer and decreased concentration in winter [36, 44, 54]. This is mainly due to higher volatilization in atmosphere and desorption from particulate phase in higher temperature [36, 55]. Contrarily, particle-bound BFRs were shown to be significantly higher in winter than in summer, which was mainly due to fuel consumption for heating activities.

Fig. 1
figure1

Distribution of BFRs in gas and particle phases in summer (a) and winter (b)

The temperature dependence of atmospheric organic pollutants can be parameterized by the Clausius–Clapeyron equation (Eq. 1) [56]. It is used to identify whether these gaseous pollutants come from long-range atmospheric transport or local surface release [19].

$$\ln P = - \frac{{\Delta H_{\text{vap}} }}{RT} + b = \frac{m}{T} + b,$$
(1)

where P is the partial pressure of the BFRs in gas phase (atm) calculated by the ideal gas state equation, ΔHvap is the energy of the phase transition by which a molecule volatilizes from the terrestrial or aquatic environment into the atmosphere (kJ/mol), R is the gas constant (8.314 J/(mol K)), T is the ambient temperature (K), while m and b are constants.

BTBPE, BDE-28, -66, -99, -100, and -154 showed significant positive dependence on temperature, accounting for 28–67% (r2) of the variations (Additional file 1: Table S8). Compared to previous researches, gaseous BFRs were positively correlated with temperature in e-waste sites in southern China (44–78%) [57] and less dependent in Lhasa (3.5–44.1%) [52], lake Chaohu, China (18–44%) [21]. The slope m of these mentioned compounds in this study (− 15,626 to − 7310.9) was much steeper than Lhasa (− 8578 to − 1649) [52], lake Chaohu, China (− 2970 to 4400) [21], northeast China (− 7255 to 1667) [16], but shallower than Southern China with the value from − 26,340 to − 16,120 [57]. Considering high levels of Penta-BDE in the sampling site, these compounds were more commanded by temperature-driven evaporation from local contamination. As HBB, BDE-47, -153, -183 had no statistic dependence on temperature with much shallower slope m, these congeners tended to be driven by regional or long-range atmospheric transport, consistent with studies in Lhasa [52], lake Chaohu [21], and rural sites in southern China [16].

In particle phase, HBB and low-brominated congeners (except Tri-BDE, Hepta-BDE) showed significant negative correlations (P < 0.01) with temperature change (Additional file 1: Table S9). Numerous high-brominated compounds also showed negative correlations (P < 0.05) with temperature, with much lower correlation coefficient than those in low-brominated ones. These were consistent with previous studies [21, 53, 58] which demonstrated that temperature is an essential factor on the release of BFRs [59]. Generally, low-brominated compounds are more likely to be affected by temperature than high-brominated congeners.

Gas–particle partitioning of BFRs

The distribution of BFRs between gas and particle phases varied considerably in summer and winter (Fig. 2). High-brominated compounds were only detected in particle phase which was mainly caused by low subcooled liquid vapor pressure (P L Ο , Pa) and high octanol–air partition coefficient (KOA), making them concentrate on particles easily [55, 60,61,62]. Low-brominated congeners distributed in both gas and particle phases and the percentage in atmosphere decreased with the increasing of the bromine number. HBB partitioned in gas phase with 95% and 26% in summer and winter, while BTBPE showed just 5% and 0%, respectively. This was in agreement with previous study which suggested that around 60% of HBB could associated with gas phase [43]. In summer, BDE-28 and BDE-66 mostly distributed in gas phase, and decreased to 61% and 7% in winter. BDE-100, -99 and -154 showed similar distribution with 84%, 82% and 69% in summer in atmosphere, and only 4%, 1%, 1% in winter, respectively. The distribution of gaseous BFRs exhibited a sharp decline in winter due to the low desorption from the particulate phase. Other researches [52, 55] also reported lower levels of BFRs in gas phase during cold period; whereas elevated gas-phase concentrations were observed when ambient temperature was higher.

Fig. 2
figure2

Distributions of BFRs between gas and particle phases during summer (a) and winter (b)

Adsorption onto the aerosol surface and absorption into the aerosol organic matter have been used to describe the gas/particle partitioning of SVOCs. The linear relationships between the logarithm of gas/particle partition coefficient (KP) and the logarithm of P L Ο or KOA can describe these two mechanisms. More details of Additional file 1: Eqs. S(2)–S(6) are listed in SI. Slope m between log KP and log P L Ο is supposed to near − 1 for reaching equilibrium partitioning of adsorption and absorption processes; indeed, it often deviated from − 1 in previous studies possibly due to thermodynamic factors or non-equilibrium conditions [63, 64]. For example, the fraction of organic matter and inorganic salts (e.g., ammonium sulfate, ammonium nitrate) is likely to affect equilibrium partitioning of BFRs [37, 65, 66]. Moreover, sampling artifacts may also result in this deviation due to the underestimation or overestimation of BFRs concentrations in gas and particle phases. Basically, gaseous BFRs from PUF may be adsorbed onto the filter; semi-volatile BFRs can also evaporate from the filter and subsequently adsorbed in PUF [57, 67]. Unfortunately, it is difficult to clarify the real bias caused by sampling artifacts since the complicated factors [67]. According to Goss and Schwarzenbach, adsorption process happened with slope m steeper than − 1 and absorption came up with slope shallower than − 0.6 [68]. The slope m of 8 low-brominated compounds was − 0.642 with correlation coefficient R2 0.6 (Fig. 3), indicating absorption is the dominant process in this study. The m value was consistent with other researches in Greece (− 0.444, − 0.4553 and − 0.5255 for industrial, urban-traffic and urban-background site) [55], south China (− 0.762 and − 0.754 for industrial and urban site) [20], and Italy (− 0.388 to − 0.228 for urban area) [37].

Fig. 3
figure3

Plots of log KP vs. log P L Ο (a) and log KP vs. log KOA (b) at the sampling site

log KP–log KOA regression line indicates that absorptive partitioning of BFRs between gas and particle phases is the dominant distribution process [61, 69]. The slope n in the present study (0.921, R2 = 0.604) deviated slightly from 1 suggested the non-equilibrium partitioning [70]. The slope n was much steeper than m value from log KP vs. log P L Ο , indicating that absorption process was the dominating sorption mechanism in studied area [71]. Compared with other researches, n value in this study was much higher than in Greece (0.5857, 0.5878 and 0.6839 for industrial, urban-traffic and urban-background site) [55] and Italy (0.31–0.51 for urban area) [37]. This was possibly due to different sources, usage pattern and partition mechanisms of those semi-volatile BFRs.

Inhalation exposure and health risk assessment for BFRs

The estimated daily intake of BFRs through inhalation in atmosphere (EDIinh, pg/kg-bw/day) and health risk assessment were calculated for children and adults during different seasons in this study.

$${\text{EDI}}_{{({\text{inh}})}} = C \times {\text{IR}} \times {\text{EF}}/{\text{BW,}}$$
(2)

where C is BFRs concentration (pg/m3) in air and PM2.5, IR is the inhalation rate (20 m3/day for adults [72] and 12 m3/day for children [73]), EF is exposure fraction and BW is average body weight (31.8 for children, 70 for adults). 100% absorption was assumed for all target BFRs.

The hazard quotient (HQ) was used to estimate the potential non-carcinogenic health risk of PBDEs. Recommended RfD values for daily oral PBDE exposure were obtained from USEPA [74] and Khan et al. [75, 76].

$${\text{HQ}} = {\text{EDI}}_{{({\text{inh}})}} /{\text{RfD}}$$
(3)

It is noted that particle size could affect exposure dose through inhalation. Nevertheless, there were limited data for fine particles and most of the researches mentioned in this study included different sizes of particles. PM2.5 and atmospheric BFRs were used to evaluate inhalation health risk in this study. The average inhalation EDI of BFRs was found to be 3.98 pg/kg-bw/day for children, which was higher than adults as 2.15 pg/kg-bw/day (Additional file 1: Table S10, Table 2). Winter had a much higher inhalation EDI of BFRs than summer except for HBB, BDE-28, -47, -66 which tended to be absorbed in gas phase in higher temperature.

Table 2 EDI (pg/kg-bw/day) of Σ27BFRs for children and adult in different season

The HQs for inhalation exposure to 13 kinds of PBDEs in gas and PM2.5 in winter were higher than in summer except for BDE-47, -66 and -99 which were mainly detected in gas phase especially in summer (Additional file 1: Table S11). Compared with adults, children had a higher risk inhalation exposure to PBDEs. Among those congeners, BDE-47 had the highest HQ with 8.43E−07 for children and 4.56E−07 for adults, followed by BDE-99, -183 and -209. Overall, the HQs for inhalation exposure were far less than 1 which suggested that health risk for inhalation of atmospheric PBDEs was relatively low.

Conclusion

The seasonal variation, gas–particle partitioning and health risk assessment of BFRs in gas and particle phases were investigated in Shanghai in this study. The GM concentrations of Σ25BFRs were 5.96, 101.70, 49.50 and 44.21 pg/m3 in gas, TSP, PM10 and PM2.5, respectively, higher than or comparable to cities in Europe and China, but much lower than the typical e-waste site. HBB and BDE-209 were the most abundant congeners in gas and particle phases, respectively. Gaseous BFRs in summer was much higher than in winter; while the particle phase showed opposite tendency. This is mainly due to higher volatilization or desorption of atmospheric BFRs in summer and emission of BFRs in particle phase through heating and power production in winter. Significant positive correlations were found between temperature and gaseous congeners including BTBPE, BDE-28, -66, -99, -100, -154, suggesting that they were more derived from local contamination evaporation. High-brominated congeners were only absorbed in particle phase; while low-brominated compounds were distributed in both gas and particle phases. The calculation of gas/particle partitioning suggested that absorption into the aerosol organic matter for BFRs was the main mechanism in this study. For most PBDEs, the health risk assessment through inhalation was higher in winter than in summer. Future work should focus on the source apportionment with a longer monitoring time for BFRs.

Availability of data and materials

Not applicable.

Abbreviations

BFRs:

brominated flame retardants

EDI:

estimated daily intake

GC/MS:

gas chromatography/mass spectrometry

GM:

geometric mean

GSD:

geometric standard deviation

HQ:

hazard quotient

K OA :

octanol–air partition coefficient

LOD:

limit of detection

LOQ:

limit of quantification

NBFRs:

novel brominated flame retardants

NCI:

negative chemical ionization

PBDEs:

polybrominated diphenyl ethers

PM10 :

particulate matter with particle size below 10 µm

PM2.5 :

particulate matter with particle size below 2.5 µm

POPs:

persistent organic pollutants

PUF:

polyurethane foam

P L Ο :

subcooled liquid vapor pressure

QFF:

quartz fiber filter

RfD:

reference dose

SIM:

selected ion monitoring

S/N:

signal–noise ratio

SRM:

standard reference material

SVOCs:

semi-volatile organic compounds

TSP:

total suspended particulate

BDE-28:

2,4,4ʹ-tribromodiphenyl ether

BDE-47:

2,2ʹ,4,4ʹ-tetrabromodiphenyl ether

BDE-66:

2,3ʹ,4,4ʹ-tetrabromodiphenyl ether

BDE-85:

2,2ʹ,3,4,4ʹ-pentabromodiphenyl ether

BDE-99:

2,2ʹ,4,4ʹ,5-pentabromodiphenyl ether

BDE-100:

2,2ʹ,4,4ʹ,6-pentabromodiphenyl ether

BDE-153:

2,2ʹ,4,4ʹ,5,5ʹ-hexabromodiphenyl ether

BDE-154:

2,2ʹ,4,4ʹ,5,6ʹ-hexabromodiphenyl ether

BDE-183:

2,2ʹ,3,4,4ʹ,5ʹ,6-heptabromodiphenyl ether

BDE-194:

2,3,4,5,2ʹ,3ʹ,4ʹ,5ʹ-octabromodiphenyl ether

BDE-206:

2,2ʹ,3,3ʹ,4,4ʹ,5ʹ,6-nonabromodiphenyl ether

BDE-207:

2,2ʹ,3,3ʹ,4,4ʹ,5,6,6ʹ-nonabromodiphenyl ether

BDE-208:

2,2ʹ,3,3ʹ,4,5,5ʹ,6,6ʹ-nonabromodiphenyl ether

BDE-209:

decabromodiphenyl ether

BDE-77:

3,3ʹ,4,4ʹ-tetrabromodiphenyl ether

BDE-118:

2,3ʹ,4,4ʹ,5-pentabromodiphenyl ether

BDE-138:

2,2ʹ,3,4,4ʹ,5-hexabromodiphenyl ether

BTBPE:

1,2-bis(2,4,6-tribromophenoxy)ethane

HBB:

hexabromobenzene

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Acknowledgements

We would like to thank assistants in Jiaxing Tongji Research Institute of Environment during the chemical and instrumental analysis. Financial support from the Chemstrres project funded by Swedish Research Council VR (Dnr. 639-2013-6913) is gratefully acknowledged.

Funding

This work was funded by the National Natural Science Foundation of China (No. 21777124) and Swedish Research Council (Dnr. 639-2013-6913).

Author information

YM performed the chemical and data analysis, and wrote the manuscript. HZ managed part of the chemical and data analysis. YQ made the study design, revised the manuscript and is the PI of the NSFC project. XD and DN involved in study design and the revision of the manuscript. ZL provided the high volume air samplers and instructed on sampling techniques. ZZ, JZ and DY gave suggestions on experimental design. All authors read and approved the final manuscript.

Correspondence to Yanling Qiu.

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Supplementary information

Additional file 1: Table S1. Description of the target chemicals, the monitoring ion (m/z) and LOD/LOQ. Table S2. Concentrations (μg/m3) of PM10 and PM2.5 in different countries or cities during summer and winter. Table S3. Concentrations (pg/m3) of BFRs in gas and particle phase between different countries or cities. Table S4. Correlation coefficient among BFR homologues in gas phase. Table S5. Correlation coefficient among BFR homologues in TSP. Table S6. Correlation coefficient among BFR homologues in PM10. Table S7. Correlation coefficient among BFR homologues in PM2.5. Table S8. Summary of Clausius-Clapeyron Regression slopes, intercepts, coefficient r2, P-values, calculated phase transition energy (ΔHcalvap) and the phase transition energy (ΔHvap) in the idea gas state (HBB, BTBPE, BDE-28, -47, -66, -99, -100, -153, -154, and BDE-183). Table S9. Correlation coefficient between temperature and BFRs in particle phase. Table S10. The estimated daily inhalation of BFRs (EDIinh) (pg/kg-bw/day) in atmosphere for children and adults. Table S11. HQs for PBDEs in children and adults via inhalation (Gas + PM2.5). Figure S1. Daily mass concentrations of TSP, PM10 and PM2.5 during entire sampling period. Figure S2. Temperature tendency for different size of particles. Figure S3. Compositional profiles of BFRs in gas, TSP, PM10 and PM2.5 cases. Figure S4. Seasonal tendency of BFR congeners in gas and particle phase.

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Meng, Y., Zhang, H., Qiu, Y. et al. Seasonal distribution, gas–particle partitioning and inhalation exposure of brominated flame retardants (BFRs) in gas and particle phases. Environ Sci Eur 31, 61 (2019) doi:10.1186/s12302-019-0257-2

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Keywords

  • Brominated flame retardants
  • Seasonal variation
  • Gas/particle partitioning
  • Inhalation exposure